Levels and distribution of polybrominated diphenyl ethers (PBDEs) in marine fishes from Chinese coastal waters

Chemosphere 82 (2011) 18–24

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Levels and distribution of polybrominated diphenyl ethers (PBDEs) in marine fishes from Chinese coastal waters Chonghuan Xia a,b,c, James C.W. Lam d, Xiaoguo Wu a,b,c, Liguang Sun a, Zhouqing Xie a,⇑, Paul K.S. Lam b,⇑ a

Institute of Polar Environment, School of Earth and Space Sciences, University of Science and Technology of China, Hefei, Anhui 230026, China State Key Laboratory in Marine Pollution and Department of Biology and Chemistry, City University of Hong Kong, Kowloon, Hong Kong, China c Advanced Laboratory for Environmental Research and Technology (ALERT), USTC-CityU Joint Advanced Research Center, Suzhou 215123, China d School of Biological Sciences, The University of Hong Kong, Pokfulam, Hong Kong, China b

a r t i c l e

i n f o

Article history: Received 29 April 2010 Received in revised form 5 September 2010 Accepted 9 October 2010 Available online 3 November 2010 Keywords: Polybrominated diphenyl ethers (PBDEs) Yellow croakers Silver pomfrets Fishes China Coast

a b s t r a c t Concentrations of polybrominated diphenyl ethers (PBDEs) in yellow croakers (Pseudosciaena crocea) and silver pomfrets (Pampus argenteus) collected from nine coastal cities along the eastern China coastline were investigated. PBDE congeners with mono- to hexa-brominated substitutions were detected in the samples, indicating their ubiquitous distribution in the marine environment of China. The total PBDE concentration averaged 3.04 ng g1 lipid wt, a level that was relatively lower than in other regions of the world, especially North America where Penta-BDE was extensively used. Geographically, the highest concentration of PBDEs was found in Xiamen, and the PBDE levels in yellow croakers were significantly higher than those in pomfrets in most of the selected cities, a pattern which may be related to the different feeding habits of the two species. The congener profiles of PBDEs were found to be different from the commonly detected pattern in fishes from other regions of the world (i.e., BDE47 > BDE99, BDE100 > BDE153, BDE154). BDE47 and BDE154 were the predominant congeners in both species, accounting for more than 60% of the total PBDE concentrations. The reasons for the relatively high proportion of BDE154 may be due to the debromination of higher brominated congeners such as BDE183 and BDE209 by these two species. Ó 2010 Elsevier Ltd. All rights reserved.

1. Introduction Polybrominated diphenyl ethers (PBDEs) have been widely used as flame retardant additives in electronic equipment, plastics, textiles, building materials, carpets, and in vehicles and aircraft (Sellstrom et al., 1993; Allchin et al., 1999). The worldwide market demand for PBDEs increased from 40 000 t in 1992 (Hale et al., 2001) to approximately 67 000 t in 2001 (Voorspoels et al., 2003) (t = metric ton). Three major commercial PBDE mixtures have been produced with varying degrees of bromine substitution on the aromatic rings. They are the Deca-product, which consists of 97–98% decabromodiphenyl ether (Deca-BDE); the Octa-product, which consists of 10–12% hexabromodiphenyl ethers (hexa-BDEs), 43– 44% heptabromodiphenyl ethers (Hepta-BDEs), and 31–35% octabromodiphenyl ethers (Octa-BDEs); and the Penta-product, which consists of 50–62% pentabromodiphenyl ethers (Penta-BDEs) and 24–38% tetrabromodiphenyl ethers (Tetra-BDEs) (Darnerud et al., 2001).

⇑ Corresponding authors. Tel./fax: +86 551 3601415 (Z. Xie); tel.: +852 2788 7681 (P.K.S. Lam). E-mail addresses: [email protected] (Z. Xie), [email protected] (P.K.S. Lam). 0045-6535/$ - see front matter Ó 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2010.10.037

PBDEs can enter the environment in several ways: (i) leaching during production and application processes; (ii) volatilization and leaching during use; (iii) particulate losses during use and disposal; and (iv) releases from PBDE-containing waste which is either incinerated as municipal waste or deposited in landfills (Darnerud et al., 2001; Voorspoels et al., 2003). PBDEs are of environmental concerns due to their persistence, potential bioaccumulation, widespread distribution via atmospheric transport, and possible adverse effects on wildlife and humans (McDonald, 2002; Mai et al., 2005). Consequently, the PBDEs comprising the Penta- and Octa-products were listed as persistent organic pollutants (POPs) and their usage was restricted under the Stockholm Convention in 2009, and Deca-BDE production and use were banned by EU starting on 1 July 2008 (de Wit et al., 2010). Because of their ubiquitous distribution, high lipophilicity and inert characteristics, PBDEs have been found in a wide variety of environmental media, including air (Wang et al., 2005; Lacovidou et al., 2009), water (Luo et al., 2008; Fontana et al., 2009), soil (Zou et al., 2007), sediment (Wang et al., 2009; Yang et al., 2009), fish (Cheung et al., 2008; Wu et al., 2009; Xu et al., 2009), birds (Naert et al., 2007; Lam et al., 2008), marine mammals (Tanabe, 2007), and humans (Zhu et al., 2009; Daniels et al., 2010). Aquatic organisms are especially efficient at accumulating halogenated POPs in their bodies

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C. Xia et al. / Chemosphere 82 (2011) 18–24

from water and sediment, in additional to dietary uptake, and thus they are good bio-indicators of environmental pollution (Hale et al., 2001; Wang et al., 2007). For example, skipjack tuna (Katsuwonus pelamis) has been used as a bio-indicator to elucidate the global distribution of PBDEs (Ueno et al., 2004). In Asia, the estimated market demand for the Penta-, Octa- and Deca-BDE products in 1999 were 0 t, 2000 t and 23 000 t, respectively, contributing 0%, 52% and 42% of the global demand, while North America contributed 98%, 36% and 44%, and Europe contributed 2%, 12% and 14%, respectively, of the world demand (de Wit, 2002). Since Penta-BDE accumulates to the greatest extent in aquatic organisms, on the whole, PBDE levels in biological samples collected in East Asia are comparable to or lower than those in Europe, and lower than those in North America (Wang et al., 2007). In the past, most studies on PBDE data in abiotic and biotic environmental samples from China have focused on highly industrialized regions, especially those in the southeast. Data on the geographical distributions of PBDEs across China are still limited. The eastern coastal cities of China are characterized by intensive industrial activities and urban development, and therefore it is conceivable that PBDE contamination along the Chinese coastline may be serious, and that residents may be exposed to these toxic contaminants via seafood consumption. In this work, PBDE levels in consumer fishes (i.e., large yellow croakers (Pseudosciaena crocea) and silver pomfrets (Pampus argenteus)) from nine coastal cities in eastern China were investigated. The objectives of this study are: (i) to measure PBDE concentrations in fish samples from the marine environment of eastern China; (ii) to investigate the spatial distribution of PBDEs in fishes from the Chinese coast; and (iii) to identify the potential sources of PBDEs in these fishes. 2. Materials and methods 2.1. Sampling Nine Chinese coastal cities from north to south, namely Dalian, Tianjin, Qingdao, Shanghai (Chongmingdao), Zhoushan, Wenzhou, Fuzhou, Quanzhou and Xiamen, were selected (Fig. 1). A total of

44 wild marine fish, including 20 yellow croakers (P. crocea) and 24 silver pomfrets (P. argenteus), were randomly collected from local fish markets or fishing boats in each city between November and December 2008. Procedures for the preparation of fishes have been described previously (Jiang et al., 2005). Briefly, fish samples were wrapped with aluminum foil, kept in polyethylene bags, and kept on ice during transportation. In the laboratory, they were stored frozen at 20 °C until chemical analysis. 2.2. Materials Standards used in this study were purchased from Wellington Laboratories (Guelph, Ontario, Canada). All solvents used throughout the study were of either HPLC or pesticide grade. 2.3. Sample preparation and extraction Analysis of PBDEs was accomplished by use of previously established methods with some modifications (Hale et al., 2008; Lam et al., 2008). Each fish sample was freeze-dried, and approximately 20 g of wet weight flesh tissue (without bone) was homogenized with anhydrous Na2SO4. The mixture was then extracted by accelerated solvent extraction (ASE 200, DIONEX Inc.) using a mixture of dichloromethane (DCM) and hexane (4:1 v/v, 40 mL) at 1500 psi and 110 °C for two cycles. The extract was concentrated to 5 mL and an aliquot of 0.5 mL was used for gravimetric determination of the lipid content. Lipid was removed from the extract by using a gel permeation chromatography column (GPC; Bio-Beads S-X3, Bio-Rad Laboratories, Hercules, CA). A mixture of dichloromethane in hexane (1:1 v/v) at a flow rate of 5 mL min1 was used for elution. Before being subjected to GPC, 5 ng of each 13C12-labeled standard (13C12-labeled BDE3, BDE15, BDE28, BDE47, BDE99, BDE153, BDE154, BDE183, BDE197, BDE207, and BDE209) was added to the extract. The concentrated extract was then transferred and further purified by elution with dichloromethane and hexane (1:9 v/ v) through activated silica gel (60 Å average pore size) column. 13 C12-labeled BDE139 was added as the recovery spike and the volume was further reduced to 60 lL prior to GC analysis.

Dalian

Tianjin

Qingdao

CHINA

yellow croakerpomfret 1ng g-1 lipid wt

Shanghai

Zhoushan Wenzhou Fuzhou Quanzhou Xiamen PRD* Fig. 1. Map showing sampling locations of marine fishes collected from cities in eastern China. [PBDE concentration in yellow croakers from PRD region (Yu et al., 2009).]

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C. Xia et al. / Chemosphere 82 (2011) 18–24

Table 1 Mean concentrations of PBDEs and ratios of BDE99/BDE100 in fishes (SP- silver pomfret; LYC-large yellow croaker) from eastern China (ng g1, lipid wt). Species

N

Lipid (%)

BDE3

BDE15

BDE28

BDE47

BDE99

BDE100

BDE153

BDE154

BDE183

RPBDEs

BDE99/ BDE100

SP

3

9.76 ± 2.75a 12.84 ± 2.85 8.46 ± 2.61

3

7.42 ± 2.47

ND

0.01 (0.4) ND

2.02 ± 0.29

50:50

LYC

3

15.60 ± 1.66

3.94 ± 0.87

52:48

SP

3

3.35 ± 1.10

0.32 (8.2) ND

2.40 ± 0.37

72:28

LYC

3

25.46 ± 2.02

3.19 ± 0.53

53:47

Zhoushan

SP

2

4.04 ± 1.90

2.35 ± 0.30

/

Wenzhou

SP

3

6.31 ± 2.57

2.37 ± 0.54

40:60

LYC

3

3.62 ± 1.07

2.84 ± 2.67

49:51

SP

2

4.89 ± 0.04

2.72 ± 0.35

/

LYC

3

14.26 ± 1.7

3.78 ± 0.57

49:51

SP

2

10.81 ± 1.15

1.11 ± 0.20

58:42

LYC

3

20.24 ± 5.98

3.25 ± 0.93

47:53

SP

3

6.56 ± 2.6

5.01 ± 1.97

63:37

LYC

2

13.49 ± 0.67

5.28 ± 0.08

46:54

44

10.64

0.10 (6.4) 0.61 (13.9) 0.56 (25.5) 0.29 (14.6) 0.51 (13.0) 0.12 (5.1) 0.48 (15.1) 0.21 (9.0) 0.37 (15.8) 0.39 (13.7) 0.03 (1.2) 0.67 (17.6) 0.29 (26.1) 0.58 (17.8) 0.20 (4.0) 0.91 (17.2) 0.40 (13.1)

36:64

SP

0.17 (11.2) 0.19 (4.4) 0.17 (7.8) 0.12 (5.9) 0.15 (3.9) 0.08 (3.1) 0.16 (4.9) 0.06 (2.7) 0.10 (4.3) 0.08 (2.8) 0.19 (7.0) 0.20 (5.2) 0.15 (13.5) 0.18 (5.5) 0.23 (4.6) 0.26 (4.8) 0.15 (5.1)

2.21 ± 0.31

Qingdao

0.01 (0.4) 0.33 (7.5) 0.20 (8.9) 0.12 (6.0) 0.32 (8.2) 0.03 (1.4) 0.27 (8.3) ND

52:48

3

0.03 (2.2) 0.36 (8.1) 0.11 (5.0) 0.12 (6.0) 0.34 (8.7) 0.09 (3.7) 0.31 (9.6) 0.12 (5.2) 0.09 (3.9) 0.21 (7.5) 0.05 (1.8) 0.33 (8.7) 0.07 (6.4) 0.27 (8.2) 0.13 (2.6) 0.41 (7.8) 0.19 (6.3)

4.41 ± 0.80

SP

1.16 (75.5) 2.02 (45.9) 1.03 (46.4) 1.28 (63.2) 1.90 (48.3) 1.90 (79.3) 1.57 (49.1) 1.62 (69.0) 1.59 (67.1) 1.26 (44.5) 2.31 (85.0) 1.90 (50.3) 0.47 (42.8) 1.71 (52.7) 4.29 (85.7) 2.60 (49.2) 1.79 (59.0) 7

0.03 (1.7) ND

Tianjin

0.02 (1.5) 0.14 (3.3) 0.05 (2.3) 0.04 (2.2) 0.17 (4.3) 0.04 (1.7) 0.20 (6.3) 0.04 (1.7) 0.05 (2.1) 0.06 (2.0) ND

85:15

3

0.01 (0.6) 0.14 (3.2) 0.08 (3.8) 0.04 (2.0) 0.21 (5.3) 0.13 (5.6) 0.05 (1.6) 0.10 (4.3) 0.02 (0.8) 0.03 (1.1) 0.05 (1.7) 0.03 (0.9) 0.04 (3.8) 0.05 (1.6) 0.03 (0.6) 0.06 (1.1) 0.07 (2.2)

1.53 ± 0.38

LYC

0.01 (0.7)b 0.60 (13.6) NDc

3.04

51:49

Dalian

Shanghai

Fuzhou

Quanzhou

Xiamen

Mean PRDd a b c d e

LYC

13

0.16 (4.9) 0.19 (8.2) ND 0.59 (20.6) 0.07 (2.7) 0.29 (7.6) ND 0.01 (0.3) ND 0.44 (8.3) 0.17 (5.5)

0.02 (0.7) 0.02 (1.7) 0.13 (3.9) 0.04 (0.8) 0.10 (1.8) 0.07 (2.4)

0.14 (5.7) 0.22 (7.7) ND 0.34 (9.0) 0.05 (4.7) 0.31 (9.5) 0.08 (1.5) 0.49 (9.2) 0.18 (6.1)

0.01 (0.3) ND 0.01 (0.2) ND 0.01 (0.2) 0.01 (0.2) 0.01 (0.5) 0.01 (0.2) 0.01 (0.7) 0.01 (0.3) 0.01 (0.2) 0.02 (0.4) 0.01 (0.3)

13e

Arithmetic mean ± standard error. Values in parentheses represent the percent contribution of particular BDE congener to RPBDEs in the fish sample. ND = not detected, and assumed as 0 for the calculation of total PBDE values. Data from Yu et al. (2009), median value. RPBDEs were reported as the sum of BDE28, BDE47, BDE66, BDE99, BDE100, BDE153, and BDE154.

2.4. Instrumental analysis Quantification of PBDEs was carried out using a GC (Agilent 7890A) equipped with a mass-selective detector (Agilent 5975c) for mono- to Deca-BDEs, using electron impact (EI) mode (for more details, please refer to Lam et al. (2008)). Fourteen major PBDE congeners were quantified using the isotope dilution method to their corresponding 13C12-labeled congeners. Lower (BDE3, BDE15, BDE28, BDE47, BDE99, BDE100, BDE153, BDE154, and BDE183) and higher (BDE196, BDE197, BDE206, BDE207, and BDE209) brominated PBDEs were analyzed by 30 m DB-5 ms and 15 m DB-5HT columns, respectively.

ples, total PBDE concentrations (RPBDEs) were reported as the sum of the masses of the nine individual PBDE congeners quantified. 2.6. Statistical analysis Statistical tests were performed with SPSS software (SPSS 17.0 for Windows, SPSS Inc.). Kruskal–Wallis nonparametric tests were used to compare the PBDE concentrations among samples from the nine coastal cities, and differences between the two fish species were compared using Mann–Whitney U nonparametric tests. Statistical significance was accepted at p < 0.05.

2.5. Quality control

3. Results and discussion

Procedural blanks were analyzed simultaneously with every batch of five samples to check for interferences or contamination from solvent and glassware. Instrumental detection limits (IDLs) were estimated as the average signal of the blanks plus three times the standard deviation of the signal of the blanks. Results were reported as ‘‘not detected” (ND) when the concentration was lower than IDLs. Recovery of 13C12-labeled PBDEs ranged between 50% and 115%. The efficiencies of ASE extraction and clean-up procedures were checked prior to the chemical analysis, and the recovery rates of 13C12-labeled standards ranged between 70% and 120% (n = 5). Since Octa- to Deca-BDEs were not detected in most sam-

3.1. PBDE concentrations in fish muscles and spatial distribution Of the 14 PBDE congeners measured, nine compounds (BDE3, BDE15, BDE28, BDE47, BDE99, BDE100, BDE153, BDE154, and BDE183) were always detected (55%, 86%, 93%, 100%, 100%, 80%, 100%, 95% and 57%) in the fish samples. The occurrence frequencies for other highly brominated congeners, BDE196, BDE197, BDE206, BDE207 and BDE209, were 9%, 2%, 11%, 2% and 2%, respectively. Therefore, only data for the nine lower brominated PBDE congeners are reported herein. Table 1 summarizes the concentrations of nine PBDE congeners based on lipid-normalized results in fishes

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C. Xia et al. / Chemosphere 82 (2011) 18–24

from different regions in eastern China. RPBDEs in all fish samples averaged 3.04 ng g1 lipid wt (lw) with a range from 1.11 to 5.28 ng g1 lw. The highest RPBDE concentrations for both yellow croakers and pomfrets occurred in Xiamen. The average RPBDE level in yellow croakers measured in this study was comparable to that in croakers previously reported for Hong Kong (average: 5.36 ng g1 lw) (Cheung et al., 2008). However, it was much lower than the reported value in yellow croakers from the Pearl River Estuary, which was from 6.3 to 54 ng g1 lw with a median value of 13 ng g1 lw (Yu et al., 2009). The geographical distribution of RPBDEs in yellow croakers and pomfrets collected from Chinese coastal waters was also investigated (Fig. 1). RPBDEs in yellow croakers were significantly higher than those in pomfrets (p < 0.05) except for croakers from Wenzhou and Xiamen; however, the average RPBDE concentrations in croakers from both Wenzhou and Xiamen were higher than those in pomfrets, which may be related to the feeding habits of the two

species. PBDEs are particle-affiliated organic compounds which are prone to sink in sediments (Xu et al., 2009). Yellow croakers often prey upon benthic organisms which feed on organic particles, and they can be expected to accumulate PBDEs to a greater degree, while silver pomfrets are plankton feeders. Geographically, no significant difference in PBDE levels among the cities was observed in yellow croakers (p = 0.379), but a significant difference was found for pomfrets (p = 0.038) which was attributed to significantly higher PBDE levels in pomfrets in Xiamen than in the other cities (p < 0.05). The high PBDE levels in Xiamen may be caused by rapid industrial development and population growth since the 1980s as it is one of five ‘‘special economic zones” in China. PBDE concentrations in fishes from Zhoushan (1.32–10.3 ng g1 lw) previously investigated (Miyake et al., 2008) were comparable to the values obtained in this work, and substantially higher concentration in fishes from the Pearl River Delta (PRD) region have also been reported (Yu et al., 2009), possibly due to this region being one of

Table 2 Global PBDE levels in marine fish muscles (ng g1, lipid weight). Location

Species

RPBDE

BDE47 Mean (min  max)

Number of measured PBDE congeners

References

Mean (min  max)

2.09 (1.33– 3.28) 9.52

6

Polar cod

3.55 (2.53– 4.98) 12.6

Eastern Hudson Bay coastline Western and southern of Norwegian coast North Sea (eastern coast of the UK) Northwest Atlantic (Maine) Georgia coast, USA

Arctic cod, capelin and sculpin

(9.8–73)

(5.3–25.9)

15

Atlantic cod

(52.5–86.0)

(38.6–62.0)

7

Wolkers et al. (2004) Jenssen et al. (2007) Kelly et al. (2008) Jenssen et al. (2007)

Whiting, herring and cod

(48.7–68.9)

(26–43)

6

Silver hake, Atlantic Mackerel, white hake, Atlantic herring, alewife, American plaice and winter flounder

62 (18.2– 81.5) (10–337)

(8.3–42)

16

California coast, USA

Shiner surf perch, white surf perch, rainbow surf perch, lingcod, speckled sanddab, striped bass, white croaker, black rockfish, canary rockfish, kelp rockfish, jack smelt, opaleye, spotted sand bass, yellowfin croaker, and Pacific mackerel Striped mullet, red drum, spotted seatrout, hardhead catfish and silver perch

302.17 (13.25– 1023.75)

162.74 (0.22– 687.50)

5

(8.0–87.5)

(3.6–31.8)

13

Mullet, bream, flathead, tailor and longtom

(5.5–37)

7

Flounder, tailor, yellowfin bream, luderick, fanbelly leatherjacket and sea mullet Yellow croaker and silver pomfret

(24.0–115)

(3.10– 27.00) (13.2–78.2)

3.04 (1.11– 5.28)

1.79 (0.47– 4.29)

9

Storfjorden, eastern Svalbard Bear island

Atlantic coast of Florida Queensland, Australia Sydney Harbour, Australia Eastern China coastline

Polar cod

Silver pomfret BDE3

BDE15

7

Boon et al. (2002) Shaw et al. (2009) Sajwan et al. (2008) Brown et al. (2006)

JohnsonRestrepo et al. (2005) Hermanussen et al. (2008) Losada et al. (2009) This study

9

Large yellow croaker BDE28

BDE47

BDE99

BDE100

BDE153

BDE154

BDE183

Xiamen Quanzhou Fuzhou Wenzhou Zhoushan Shanghai Qingdao Tianjin Dalian 0%

20%

40%

60%

80%

100%

0%

20%

40%

60%

Fig. 2. PBDE congener profiles in silver pomfrets (Pampus argenteus) and yellow croakers (Pseudosciaena crocea).

80%

100%

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the most economically developed areas in China with intense electronics, plastics, and textile manufacturing activities. Many studies have also reported relatively higher levels of PBDEs in the PRD region in sediments (Mai et al., 2005), water (Luo et al., 2008) and fishes (Guo et al., 2008) compared to other regions in China. Combining data reported for the PRD region with the results of the present work showed that RPBDE levels in fishes along the coast of eastern China were higher in the southern than in the central and the northern coastal areas. Table 2 shows a global view of the PBDE levels in marine fish. RPBDE concentrations and concentrations of an individual congener (BDE47, commonly the most abundant congener in biological samples) are shown. Results of RPBDE and BDE47 concentrations were comparable to those in polar cods from eastern Svalbard. Overall, the concentrations of RPBDE and BDE47 in marine fish in the present study were relatively lower when compared with the other regions of the world, especially North America. This is in agreement with other studies (Meng et al., 2007; Wan et al., 2008; Xu et al., 2009), although global monitoring of PBDEs using skipjack tuna as a bio-indicator and the Asian mussel watch program both reported relatively high concentrations of PBDEs in China (Ueno et al., 2004; Ramu et al., 2007). The main reason for this trend is most likely due to the heavier usage of Penta-BDE in North America than in Europe and Asia (Hites, 2004). In Asia, the consumption of Penta-BDE (150 t) contributed only 2.0% of the global demand in 2001, far less than the amount (7100 t) used in North America (Law et al., 2006). In addition, the congeners in the Penta-BDE mixture are highly bioaccumulative. Metabolic differences, age of the fish, and lipid content as well as other physiological differences among fish species are also possible reasons for the observed PBDE patterns (Haglund et al., 1997). For example, Yu et al. (2009) reported PBDE levels in fishes from the PRD, and found that yellow croakers had relatively lower PBDE concentrations than other species, such as pompano (Psenopsis anomala), flathead fish (Platycephalus indicus) and robust tonguefish (Cynoglossus robustus), even though these species occupied comparable trophic positions. 3.2. Congener profiles of PBDEs Contributions of each PBDE congener to RPBDEs in the two species of fishes are presented in Fig. 2. The most abundant congener was BDE47 with the percentage ranging between 43% and 86% of RPBDE concentrations. BDE47 has been previously found to be the predominant congener in fish (Peng et al., 2007; Xian et al., 2008; Xu et al., 2009). Congener concentrations in fish from high to low were generally as follows: BDE47 > BDE99, BDE100 > BDE153, BDE154 (Boon et al., 2002; Hites, 2004; Xiang et al., 2007). For example, in fishes from Europe, these five congeners accounted for 69.1%, 15.4%, 12.8%, 2.8% and 4.8% of RPBDEs concentrations, respectively (Hites, 2004). However, overall PBDE congener profiles in fishes in this study differed from those reported previously. In this work, BDE47, BDE99, BDE100, BDE153, BDE154 and BDE3 accounted for 68.2%, 4.1%, 3.2%, 6.7%, 12.0% and 1.3% of the RPBDE concentration in pomfrets and 47.8%, 8.3%, 8.3%, 4.5%, 15.2% and 10.4% of the RPBDE concentration in yellow croakers, respectively. Notably, BDE154 was found at relatively higher proportions than BDE99 and BDE100 in both species, a pattern which was also found in seafood from China (Shen et al., 2009). A study on PBDEs in fish samples in Taiwan also indicated that BDE154 contributed more to RPBDEs than BDE99 and BDE100, which was suggested to be due to the extensive use of Octa-BDE rather than Penta-BDE (Peng et al., 2007). The relatively high BDE154 percentage found in the present study may reflect the metabolic ability of fishes. Fish may metabolize and debrominate highly brominated congeners such as BDE183 (Stapleton et al.,

2004b) and BDE209 (Stapleton et al., 2004a) to BDE154, even though BDE209 has very low bioavailability. Results of two independent dietary exposure studies using the common carp (Cyprinus carpio) to trace the fate of BDE209 and BDE183 in fish tissues indicated that a limited amount of BDE209 was bioavailable from food in the form of lower brominated metabolites, including BDE154 (Stapleton et al., 2004a); however, significant debromination of BDE183 to BDE154 was observed (Stapleton et al., 2004b). Furthermore, Deca-BDE made up primarily of BDE209 and Octa-BDE containing as much as 44% BDE183 have been used in large quantities in China; the Asian demand for Octa-BDE (2000 t) and Deca-BDE (23 000 t) accounted for 52% and 42% of the global market in 1999 (de Wit, 2002). The high proportion of BDE154 may partially explain the relatively low BDE183 percentage and non-detection of BDE209 in fishes. It is interesting that lower brominated congeners such as BDE3, which was not detected in other studies (Ueno et al., 2004; Luo et al., 2007), contributed more to RPBDEs than higher brominated congener in yellow croakers, but the frequency of detection of BDE3 was only 29% in pomfrets. Several other studies found relatively greater abundance of lower brominated congeners including BDE15 and BDE28 in sediments (Chen et al., 2006) and human breast milk (Sudaryanto et al., 2007) in China, and suggested that the usage of additional specific technical mixtures of PBDEs in China was the probable reason. For example, the Tetra-BDE product was used in Japan until 1990, but now it is no longer commercially produced (Akutsu et al., 2001). It is possible that commercial mixtures containing these lower brominated BDEs were used in China, resulting in higher residual levels of lower BDEs in the environment (Xian et al., 2008). BDE15 has been reported to photochemically degrade to BDE3 in an aqueous (H2O:CH3CN; 1:1 v/v) solvent system at 300 nm via homolytic C–Br bond cleavage (Rayne et al., 2003). This may imply that BDE15 is degrading to BDE3 in this aquatic system, increasing the likelihood that organisms will be exposed to BDE3. However, the higher concentrations and occurrence frequencies of BDE3 in yellow croakers than pomfrets may indicate the greater capability of croakers to bioaccumulate BDE3; previous studies which investigated PBDE levels in this species of yellow croakers (e.g., Shen et al., 2009; Yu et al., 2009) did not measure BDE3. Another noteworthy finding is that BDE209 was detected in only one sample, considering that the Deca-BDE product has been the main brominated fire retardant used in China. Non-detection of BDE209 in fish muscle from China has been previously reported (Meng et al., 2007; Xian et al., 2008). Because of the higher molecular weight and larger octanol/water partition coefficient (log Kow  10) of BDE209, it has lower bioavailability compared to other PBDE congeners, which may be a possible explanation for this result as suggested by Meng et al. (2007). On the other hand, debromination of BDE209 to more bioavailable metabolites has been found to occur in fish tissues (Stapleton et al., 2004a), which was discussed in detail above. However, the uptake and bioaccumulation pathways of BDE209 in fish are still not well understood. 3.3. Ratio of BDE99/BDE100 The ratios between PBDE congeners can be used to understand their mechanisms of accumulation in organisms (Borghesi et al., 2009), and there have been several interesting observations concerning the concentration ratios between BDE99 and BDE100. Christensen et al. (2002) reviewed previous investigations and found that the ratio of BDE99 and BDE100 in the abiotic environment (air and sediment) was 80:20, which was similar to the ratio in the industrial product Bromkal 70-5DE (approximately 84:16), and in biotic samples (fish and marine mammals) the ratio aver-

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aged 30:70. The possible explanations for these ratios were suggested to be the higher bioavailability of BDE100 compared to BDE99 or that BDE100 was less biodegradable compared to BDE99 (Christensen et al., 2002). In the present study, the ratios between BDE99 and BDE100 in large yellow croakers and pomfrets varied among regions. In yellow croakers, the ratios ranged from 46:54 in Xiamen to 53:47 in Shanghai, while those in pomfrets ranged from 36:64 in Tianjin to 85:15 in Dalian, confirming that the BDE99/BDE100 ratio is location-dependent (Voorspoels et al., 2003). In the same region, the ratios in the two species of fishes were also different. For example, in Dalian, the ratio in yellow croakers was 52:48, but in pomfrets it was 85:15. Voorspoels et al. (2003) reported that the ratios of BDE99/BDE100 in marine species from the Belgian North Sea ranged from 80:20 in shrimp (Crangon crangon) to 20:80 in sole (Solea solea). Another study on various organisms (three invertebrate species and five fish species) in the Pearl River estuary in south China found that the ratios varied among species from 80:20 to 25:75 (Xiang et al., 2007). These differences may be explained by the species-dependent metabolic abilities of the sampled organisms (Voorspoels et al., 2003). 4. Conclusion This study investigated PBDE levels in marine fishes collected along the eastern coastline of China. PBDEs were detected in all samples, and the highest concentration of RPBDEs was found in Xiamen. RPBDEs in yellow croakers were significantly higher than those in pomfrets in most of the selected cities, a pattern which may be related to the different feeding habits of the two species. Overall, RPBDE concentrations in fishes were relatively lower in comparison with those reported for other regions of the world, especially North America where Penta-BDE were more heavily used. A relatively high proportion of BDE154 was found, perhaps due to the debromination of higher brominated congeners, such as BDE183 and BDE209. Acknowledgements This study was partially supported by Grants from National Basic Research Program of China (973 Program 2010CB428902 and 2009CB421605), the National Natural Science Foundation of China (Project’s No. 41025020, 40776001), the Foundation for the Author of National Excellent Doctoral Dissertation of PR China (Grant 200354), the Chinese Academy of Sciences (Grant KZCX2-YWQN506) and the Fundamental Research Funds for the Central Universities. The work described in this paper was also funded by the Area of Excellence Scheme under the University Grants Committee of the Hong Kong Special Administrative Region, China (Project No. AoE/P04/2004), and a Hong Kong Research Grants Council (CityU 160610). Reference Akutsu, K., Obana, H., Okihashi, M., Kitagawa, M., Nakazawa, H., Matsuki, Y., Makino, T., Oda, H., Hori, S., 2001. GC/MS analysis of polybrominated diphenyl ethers in fish collected from the Inland Sea of Seto, Japan. Chemosphere 44, 1325–1333. Allchin, C.R., Law, R.J., Morris, S., 1999. Polybrominated diphenylethers in sediments and biota downstream of potential sources in the UK. Environmental Pollution 105, 197–207. Boon, J.P., Lewis, W.E., Tjoen-A-Choy, M.R., Allchin, C.R., Law, R.J., de Boer, J., ten Hallers-Tjabbes, C.C., Zegers, B.N., 2002. Levels of polybrominated diphenyl ether (PBDE) flame retardants in animals representing different trophic levels of the North Sea food web. Environmental Science & Technology 36, 4025–4032. Borghesi, N., Corsolini, S., Leonards, P., Brandsma, S., de Boer, J., Focardi, S., 2009. Polybrominated diphenyl ether contamination levels in fish from the Antarctic and the Mediterranean Sea. Chemosphere 77, 693–698. Brown, F.R., Winkler, J., Visita, P., Dhaliwal, J., Petreas, M., 2006. Levels of PBDEs, PCDDs, PCDFs, and coplanar PCBs in edible fish from California coastal waters. Chemosphere 64, 276–286. Chen, S.J., Gao, X.J., Mai, B.X., Chen, Z.M., Luo, X.J., Sheng, G.Y., Fu, J.M., Zeng, E.Y., 2006. Polybrominated diphenyl ethers in surface sediments of the Yangtze

23

River Delta: levels, distribution and potential hydrodynamic influence. Environmental Pollution 144, 951–957. Cheung, K.C., Zheng, J.S., Leung, H.M., Wong, M.H., 2008. Exposure to polybrominated diphenyl ethers associated with consumption of marine and freshwater fish in Hong Kong. Chemosphere 70, 1707–1720. Christensen, J.H., Glasius, M., Pecseli, M., Platz, J., Pritzl, G., 2002. Polybrominated diphenyl ethers (PBDEs) in marine fish and blue mussels from southern Greenland. Chemosphere 47, 631–638. Daniels, J.L., Pan, I.J., Jones, R., Anderson, S., Patterson, D.G., Needham, L.L., Sjodin, A., 2010. Individual characteristics associated with PBDE levels in US human milk samples. Environmental Health Perspectives 118, 155–160. Darnerud, P.O., Eriksen, G.S., Johannesson, T., Larsen, P.B., Viluksela, M., 2001. Polybrominated diphenyl ethers: occurrence, dietary exposure, and toxicology. Environmental Health Perspectives 109, 49–68. de Wit, C.A., 2002. An overview of brominated flame retardants in the environment. Chemosphere 46, 583–624. de Wit, C.A., Herzke, D., Vorkamp, K., 2010. Brominated flame retardants in the Arctic environment – trends and new candidates. Science of the Total Environment 408, 2885–2918. Fontana, A.R., Silva, M.F., Martinez, L.D., Wuilloud, R.G., Altamirano, J.C., 2009. Determination of polybrominated diphenyl ethers in water and soil samples by cloud point extraction-ultrasound-assisted back-extraction-gas chromatography–mass spectrometry. Journal of Chromatography A 1216, 4339–4346. Guo, L.L., Qiu, Y.W., Zhang, G., Zheng, G.J., Lam, P.K.S., Li, X.D., 2008. Levels and bioaccumulation of organochlorine pesticides (OCPs) and polybrominated diphenyl ethers (PBDEs) in fishes from the Pearl River estuary and Daya Bay, South China. Environmental Pollution 152, 604–611. Haglund, P.S., Zook, D.R., Buser, H.R., Hu, J.W., 1997. Identification and quantification of polybrominated diphenyl ethers and methoxypolybrominated diphenyl ethers in Baltic biota. Environmental Science & Technology 31, 3281–3287. Hale, R.C., Kim, S.L., Harvey, E., La Guardia, M.J., Mainor, T.M., Bush, E.O., Jacobs, E.M., 2008. Antarctic research bases: local sources of polybrominated diphenlyl ether (PBDE) flame retardants. Environmental Science & Technology 42, 1452–1457. Hale, R.C., La Guardia, M.J., Harvey, E.P., Mainor, T.M., Duff, W.H., Gaylor, M.O., 2001. Polybrominated diphenyl ether flame retardants in virginia freshwater fishes (USA). Environmental Science & Technology 35, 4585–4591. Hermanussen, S., Matthews, V., Paepke, O., Limpus, C.J., Gaus, C., 2008. Flame retardants (PBDEs) in marine turtles, dugongs and seafood from Queensland, Australia. Marine Pollution Bulletin 57, 409–418. Hites, R.A., 2004. Polybrominated diphenyl ethers in the environment and in people: a meta-analysis of concentrations. Environmental Science & Technology 38, 945–956. Jenssen, B.M., Sormo, E.G., Baek, K., Bytingsvik, J., Gaustad, H., Ruus, A., Skaare, J.U., 2007. Brominated flame retardants in North–East Atlantic marine ecosystems. Environmental Health Perspectives 115, 35–41. Jiang, Q.T., Lee, T.K.M., Chen, K., Wong, H.L., Zheng, J.S., Giesy, J.P., Lo, K.K.W., Yamashita, N., Lam, P.K.S., 2005. Human health risk assessment of organochlorines associated with fish consumption in a coastal city in China. Environmental Pollution 136, 155–165. Johnson-Restrepo, B., Kannan, K., Addink, R., Adams, D.H., 2005. Polybrominated diphenyl ethers and polychlorinated biphenyls in a marine foodweb of coastal Florida. Environmental Science & Technology 39, 8243–8250. Kelly, B.C., Ikonomou, M.G., Blair, J.D., Gobas, F., 2008. Bioaccumulation behaviour of polybrominated diphenyl ethers (PBDEs) in a Canadian Arctic marine food web. Science of the Total Environment 401, 60–72. Lacovidou, E., Mandalakis, M., Stephanou, E.G., 2009. Occurrence and diurnal variation of polychlorinated biphenyls and polybrominated diphenyl ethers in the background atmosphere of Eastern Mediterranean. Chemosphere 77, 1161– 1167. Lam, J.C.W., Murphy, M.B., Wang, Y., Tanabe, S., Giesy, J.P., Lam, P.K.S., 2008. Risk assessment of organohalogenated compounds in water bird eggs from South China. Environmental Science and Technology 42, 6296–6302. Law, R.J., Allchin, C.R., de Boer, J., Covaci, A., Herzke, D., Lepom, P., Morris, S., Tronczynski, J., de Wit, C.A., 2006. Levels and trends of brominated flame retardants in the European environment. Chemosphere 64, 187–208. Losada, S., Roach, A., Roosens, L., Santos, F.J., Galceran, M.T., Vetter, W., Neels, H., Covaci, A., 2009. Biomagnification of anthropogenic and naturally-produced organobrominated compounds in a marine food web from Sydney Harbour, Australia. Environment International 35, 1142–1149. Luo, Q., Cai, Z.W., Wong, M.H., 2007. Polybrominated diphenyl ethers in fish and sediment from river polluted by electronic waste. Science of the Total Environment 383, 115–127. Luo, X.J., Yu, M., Mai, B.X., Chen, S.J., 2008. Distribution and partition of polybrominated diphenyl ethers (PBDEs) in water of the Zhujiang River Estuary. Chinese Science Bulletin 53, 493–500. Mai, B.X., Chen, S.J., Luo, X.J., Chen, L.G., Yang, Q.S., Sheng, G.Y., Peng, P.G., Fu, J.M., Zeng, E.Y., 2005. Distribution of polybrominated diphenyl ethers in sediments of the Pearl River Delta and adjacent South China Sea. Environmental Science & Technology 39, 3521–3527. McDonald, T.A., 2002. A perspective on the potential health risks of PBDEs. Chemosphere 46, 745–755. Meng, X.Z., Zeng, E.Y., Yu, L.P., Guo, Y., Mai, B.X., 2007. Assessment of human exposure to polybrominated diphenyl ethers in China via fish consumption and inhalation. Environmental Science & Technology 41, 4882–4887.

24

C. Xia et al. / Chemosphere 82 (2011) 18–24

Miyake, Y., Jiang, Q., Yuan, W., Hanari, N., Okazawa, T., Wyrzykowska, B., So, M.K., Lam, P.K.S., Yamashita, N., 2008. Preliminary health risk assessment for polybrominated diphenyl ethers and polybrominated dibenzo-p-dioxins/ furans in seafood from Guangzhou and Zhoushan, China.. Marine Pollution Bulletin 57, 357–364. Naert, C., Van Peteghem, C., Kupper, J., Jenni, L., Naegeli, H., 2007. Distribution of polychlorinated biphenyls and polybrominated diphenyl ethers in birds of prey from Switzerland. Chemosphere 68, 977–987. Peng, J.H., Huang, C.W., Weng, Y.M., Yak, H.K., 2007. Determination of polybrominated diphenyl ethers (PBDEs) in fish samples from rivers and estuaries in Taiwan. Chemosphere 66, 1990–1997. Ramu, K., Kajiwara, N., Sudaryanto, A., Isobe, T., Takahashi, S., Subramanian, A., Ueno, D., Zheng, G.J., Lam, P.K.S., Takada, H., Zakaria, M.P., Viet, P.H., Prudente, M., Tana, T.S., Tanabe, S., 2007. Asian mussel watch program: contamination status of polybrominated diphenyl ethers and organochlorines in coastal waters of Asian countries. Environmental Science & Technology 41, 4580–4586. Rayne, S., Ikonomou, M.G., Whale, M.D., 2003. Anaerobic microbial and photochemical degradation of 4, 40 -dibromodiphenyl ether. Water Research 37, 551–560. Sajwan, K.S., Kumar, K.S., Nune, S., Fowler, A., Richardson, J.P., BG, L., 2008. Persistent organochlorine pesticides, polychlorinated biphenyls, polybrominated diphenyl ethers in fish from coastal waters off Savannah, GA, USA. Toxicological and Environmental Chemistry 90, 81–96. Sellstrom, U., Jansson, B., Kierkegaard, A., Dewit, C., Odsjo, T., Olsson, M., 1993. Polybrominated diphenyl ethers (PBDE) in biological samples from the Swedish environment. Chemosphere 26, 1703–1718. Shaw, S.D., Berger, M.L., Brenner, D., Kannan, K., Lohmann, N., Papke, O., 2009. Bioaccumulation of polybrominated diphenyl ethers and hexabromocyclododecane in the northwest Atlantic marine food web. Science of the Total Environment 407, 3323–3329. Shen, H.T., Yu, C., Ying, Y., Zhao, Y.F., Wu, Y.N., Han, J.L., Xu, Q.Y., 2009. Levels and congener profiles of PCDD/Fs, PCBs and PBDEs in seafood from China. Chemosphere 77, 1206–1211. Stapleton, H.M., Alaee, M., Letcher, R.J., Baker, J.E., 2004a. Debromination of the flame retardant decabromodiphenyl ether by juvenile carp (Cyprinus carpio) following dietary exposure. Environmental Science & Technology 38, 112– 119. Stapleton, H.M., Letcher, R.J., Baker, J.E., 2004b. Debromination of polybrominated diphenyl ether congeners BDE 99 and BDE 183 in the intestinal tract of the common carp (Cyprinus carpio). Environmental Science & Technology 38, 1054– 1061. Sudaryanto, A., Kajiwara, N., Tsydenova, O.V., Kunisue, T., Yu, H., Tanabe, S., 2007. Levels and congener specific profiles of PBDE in human breast milk from mothers living in Nanjing and Zhoushan, China. Chemical pollutions and environmental changes. In: Proceedings of the 4th International Symposium on Pioneering Studies of Young Scientists on Chemical Pollutions and Environmental Changes November 17–19, 2006, Universal Academy Press Inc., pp. 2171–2174. Tanabe, S., 2007. Temporal trends of brominated flame retardants in coastal waters of Japan and South China: Retrospective monitoring study using archived samples from es-Bank, Ehime University, Japan. In: 5th International Conference on Marine Pollution and Ecotoxicology, Hong Kong, Peoples R China, pp. 267–274.

Ueno, D., Kajiwara, N., Tanaka, H., Subramanian, A., Fillmann, G., Lam, P.K.S., Zheng, G.J., Muchitar, M., Razak, H., Prudente, M., Chung, K.H., Tanabe, S., 2004. Global pollution monitoring of polybrominated diphenyl ethers using skipjack tuna as a bioindicator. Environmental Science & Technology 38, 2312–2316. Voorspoels, S., Covaci, A., Schepens, P., 2003. Polybrominated diphenyl ethers in marine species from the Belgian North Sea and the Western Scheidt Estuary: levels, profiles, and distribution. Environmental Science & Technology 37, 4348–4357. Wan, Y.I., Hu, J.Y., Zhang, K., An, L.H., 2008. Trophodynamics of polybrominated diphenyl ethers in the marine food web of Bohai Bay, North China. Environmental Science & Technology 42, 1078–1083. Wang, X.M., Ding, X., Mai, B.X., Xie, Z.Q., Xiang, C.H., Sun, L.G., Sheng, G.Y., Fu, J.M., Zeng, E.Y., 2005. Polyhrominated diphenyl ethers in airborne particulates collected during a research expedition from the Bohai Sea to the Arctic. Environmental Science & Technology 39, 7803–7809. Wang, Y.W., Jiang, G.B., Lam, P.K.S., Li, A., 2007. Polybrominated diphenyl ether in the East Asian environment: a critical review. Environment International 33, 963–973. Wang, Z., Ma, X.D., Lin, Z.S., Na, G.S., Yao, Z.W., 2009. Congener specific distributions of polybrominated diphenyl ethers (PBDEs) in sediment and mussel (Mytilus edulis) of the Bo Sea, China. Chemosphere 74, 896–901. Wolkers, H., Van Bavel, B., Derocher, A.E., Wiig, O., Kovacs, K.M., Lydersen, C., Lindstrom, G., 2004. Congener-specific accumulation and food chain transfer of polybrominated diphenyl ethers in two Arctic food chains. Environmental Science & Technology 38, 1667–1674. Wu, J.P., Luo, X.J., Zhang, Y., Yu, M., Chen, S.J., Mai, B.X., Yang, Z.Y., 2009. Biomagnification of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls in a highly contaminated freshwater food web from South China. Environmental Pollution 157, 904–909. Xian, Q.M., Ramu, K., Isobe, T., Sudaryanto, A., Liu, X.H., Gao, Z.S., Takahashi, S., Yu, H.X., Tanabe, S., 2008. Levels and body distribution of polybrominated diphenyl ethers (PBDEs) and hexabromocyclododecanes (HBCDs) in freshwater fishes from the Yangtze River, China. Chemosphere 71, 268–276. Xiang, C.H., Luo, X.J., Chen, S.J., Yu, M., Mai, B.X., Zeng, E.Y., 2007. Polybrominated diphenyl ethers in biota and sediments of the Pearl River Estuary, South China. Environmental Toxicology and Chemistry 26, 616–623. Xu, J., Gao, Z.S., Xian, Q.M., Yu, H.X., Feng, J.F., 2009. Levels and distribution of polybrominated diphenyl ethers (PBDEs) in the freshwater environment surrounding a PBDE manufacturing plant in China. Environmental Pollution 157, 1911–1916. Yang, Z.Z., Zhao, X.R., Qin, Z.F., Fu, S., Li, X.H., Qin, X.F., Xu, X.B., Jin, Z.X., 2009. Polybrominated diphenyl ethers in mudsnails (Cipangopaludina cahayensis) and sediments from an electronic waste recycling region in South China. Bulletin of Environmental Contamination and Toxicology 82, 206–210. Yu, M., Luo, X.J., Wu, J.P., Chen, S.J., Mai, B.X., 2009. Bioaccumulation and trophic transfer of polybrominated diphenyl ethers (PBDEs) in biota from the Pearl River Estuary, South China. Environment International 35, 1090–1095. Zhu, L.Y., Ma, B.L., Li, J.G., Wu, Y.N., Gong, J., 2009. Distribution of polybrominated diphenyl ethers in breast milk from North China: implication of exposure pathways. Chemosphere 74, 1429–1434. Zou, M.Y., Ran, Y., Gong, J., Maw, B.X., Zeng, E.Y., 2007. Polybrominated diphenyl ethers in watershed soils of the Pearl River Delta, China: occurrence, inventory, and fate. Environmental Science & Technology 41, 8262–8267.